Download biodiversity and infectious disease: why we need nature

Interactions between Global Change and Human Health
Pontifical Academy of Sciences, Scripta Varia 106, Vatican City 2006
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BIODIVERSITY AND INFECTIOUS DISEASE:
WHY WE NEED NATURE
ANDREW P. DOBSON
Preface
The large leather bible hit the congressional committee room table
with a crash that commanded the attention of congressmen, their aides,
and the press assembled for the Endangered Species Act hearings.
‘Extinction is a sin – We are destroying God’s creation’, announced the
President of Christians for the Environment. This produced a nervous
shuffling in seats. Congressmen unmoved by economic or scientific arguments were plainly disconcerted by this sharp reminder that our moral
responsibility to other species puts parts of their constituency in blunt
conflict with each other. The rest of us breathed a little more easily. The
current US Congress, curiously immune to economic and scientific arguments, has a visceral response to the Old Testament. Here I will expand
on this testimony and outline the broader economic, scientific and ethical arguments for ‘Why we need nature’.
Biodiversity as Food-Webs
One of the first things we learn about in school nature classes is food
chains. My five-year-old son can happily arrange krill, fish, big fish, and
sharks into a logical hierarchy of producers and consumers. Ecologists and
economists have long been fascinated with the mathematical properties of
simplified food webs (Dunne, 2005). The Italian mathematician, Vito
Volterra, initially examined the properties of the simplest one predator and
one prey food web in the 1920s (Volterra, 1926). His stimulus was the worry
that his daughter would marry a fisherman. When he created a mathematical model of the interactions between fisherman and fish he realized that
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that even this simple model had a pathological tendency to undergo sustained cycles of abundance of predators (fisherman) and prey (fish). While
this was an unsuccessful deterrent to his daughter’s marriage, it catalyzed
a massive expansion in mathematical interest in the properties of ecological systems. This early work of Volterra and Lotka (Lotka, 1923) advanced
our understanding of consumer-resource relationships considerably
beyond Malthus, whose worries about the geometric and arithmetic rates
of increase of consumer and resource are but a sub-set of this broad and
complex set of problems. The mathematical study of predator-prey systems
is at the heart of many natural processes; lions chasing zebra, insects eating plants and each-other, right through to how the immune system interacts with viral and other pathogens that invade our bodies (May, 2001;
Nowak and May, 1991). These problems rapidly become more complex as
we increase the numbers of species that interact as hierarchical networks
of consumer and resource species within a food web (Cohen, 1989; Cohen
et al., 1990; Cohen et al., 2003; Dunne et al., 2002).
‘There Ain’t Have Been Some Clever Geezers…’
To understand the scientific complexity of the problem we have to
compare ecology and the other life sciences to physics and chemistry. One
way to do this is to reduce these sciences to the level of the numbers of
different particles they consider, the spatial and temporal scales at which
those particles interact, and the complexity, or non-linearity, of these
interactions. In their purest form physics and chemistry consider limited
numbers of particles interacting at either extremely small or rapid scales,
or at huge scales over rapid timescales. The sheer magnitude of these
scales creates a mystical awe. Ironically this masks their lack of utility to
any aspect of human health or well being! In contrast, biologists consider interactions between large numbers of particles of different types (molecules, cells, species) embedded in a complex hierarchy of webs, where
processes occur at timescales from seconds and minutes to millennia.
Because many of the species involved are cute, charismatic, or pests, and
studied at scales closer to fish-tanks and fields, there is less mystical awe
in the study of their interactions.
When we convert forest or savannas to agricultural land we simplify
the local food webs and focus on cultivating a handful of domesticated
species as resources. This simplification reduces our potential to benefit
from the services supplied by the natural web. How we value nature is
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147
intimately linked with understanding the trade-offs between agriculture,
trade and nature (Arrow et al., 1995; Balmford et al., 2002; Costanza,
1991). Inherently asking the question ‘Do we need Nature?’ assumes we
can realistically compare the long-term value of natural habitats, whose
benefits are diffusely appreciated by a variety of people, with the instantaneous and focused annual yields of agriculture, timber products, or golf
course revenues. Some of the problems associated with this calculation
boil down to simply not knowing if the land is worth more as agricultural land or natural habitat (Balmford et al., 2002). If we focus on a single
service or product we will always underestimate its natural value. If we
try and completely enumerate its value, we run the risk of exaggeration.
Even if we could undertake these calculations, we’d then need to quantify the dependence of the modified habitat on services produced by the
natural habitat (Dobson, 2005b).
In the absence of a comprehensive pricing of nature’s services, then
the natural habitat will appear to be worth less than when converted to
agricultural land. This creates further pressure to convert all of the land
to agriculture. As an example, consider an area of land containing a patch
of woodland or prairie – this could range in size from a small farm, to
many hundreds of square miles, the Amazon basin, or the Serengeti. If we
convert the forest or savanna to agricultural land (or a golf course), its
economic value will be determined by the quality and quantity of products and services produced, minus the cost of converting the land. In contrast, if we leave the land in an undisturbed state it will have an economic value determined by a complex calculation that sums the resources
people remove from it for food or fiber, plus the added value we obtain as
it cleanses the surrounding air and water. We may also have to add its
value as a sink for greenhouse gases. Finally we must include a more
complex value determined by the pleasure people receive from spending
time there, or even simply knowing it is there. Alternatively, we could convert a proportion of the land to agriculture and leave the rest as forest.
This may be particularly sensible if the agricultural land is dependent
upon the forest for services such as a constant water supply, or a source
of pollinators that ensure crop fertility (Figure 1, see page 407).
Determining the proportion of land to convert and the proportion to
leave available for other species is the central problem of natural resource
management. Unfortunately the less we value Nature then the smaller the
proportion of land that will be kept set aside for other species. Many
international conservation treaties recommend leaving ten percent of the
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land as set aside for Nature. Their logic stems from the well-known
‘species-area’ relationship discovered by Robert MacArthur and E.O.
Wilson in the 1960s (MacArthur and Wilson, 1963; MacArthur and
Wilson, 1967); this relationship suggests that each successive 90% loss in
habitat dooms 50% of remaining species to local extinction. From a purely utilitarian perspective, the amount of biodiversity we save is one
answer to the question ‘Will this level of biodiversity be sufficient to supply all the necessary services we are used to receiving from Nature?’
Unfortunately, the benefits of maintaining biodiversity are most
sharply appreciated in their absence! This usually occurs when we notice
a decline in air or water quality, reductions in crop pollination rates, or
increased pest and disease outbreaks. We are then stuck with the problem
of either replacing ‘Nature’s services’ with a technological fix, or of trying
to restore the natural habitat. Both are likely to be expensive; in the case
of restoration the significant current costs may not be met by future discounted potential benefits (Dobson, 2005b; Dobson et al., 1997;
Simberloff et al., 1998). Restoration also assumes we know enough about
ecosystem functioning to put food-webs back together (Bradshaw, 1983;
Bradshaw, 1984). This illustrates one of the deepest scientific ironies of
the 21st Century: while we know much about the structure of atoms and
of the universe, we have only the most rudimentary understanding of the
structure of salt-marshes, forests, or even the soil beneath our crops.
Reasons to be Careful… Part 3
When land is converted from natural habitat, different species will go
extinct at different rates (Seabloom et al., 2002; Tilman et al., 1994).
Species are lost as their habitats are converted to agricultural land, as we
exploit them for food, or as they are out-competed for space and
resources by a suite of invasive plants and animals that are the camp followers of human expansion (Crawley, 1986; Soulé, 1986). We are thus seeing a transformation from a world of tropical forests, tigers, orchids and
pandas, to one of weeds, mosquitoes, goats, and sparrows. This sequential loss of biodiversity means that the economic services will be lost at
different rates. Services that are predominantly supplied by species with
large area requirements, rare species, or species with specialist habitat
requirements, will be lost more rapidly than those mediated by species
that can persist in a handful of soil, or those able to rapidly adapt to the
suburban environment (Dobson et al., 2006; Jenkins, 2003; Kremen,
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2005). In some cases agricultural or weedy species may replace some of
the ecosystem services performed by the original native species. In general, the services provided by species at the top of the food chain will be
lost more rapidly than those provided by those at the base of the web
(Dobson et al., 2006; Loreau et al., 2001). This creates a hierarchical loss
of natural services, initially the aesthetic benefits will decline, orchids and
pandas will become rare, and tigers and other predators will disappear
locally (Figure 2, see page 408). Except in areas that are strongly dependent upon ecotourism, this will have only a limited impact on the local
economy. It will create problems, the loss of predators will cause the
species they prey upon to increase in abundance and become pests (Duffy,
2003). In turn this will lead to an increase in thorny, inedible vegetation
that encourages the now abundant deer, rabbits and insects to feed upon
crops. The increased contact of these species with humans and domestic
livestock increases the potential for disease transmission. Laws such as
the Endangered Species Act in the United States can prevent this cascade
of events. While focusing attention on individual species, the law seeks to
protect entire functioning food webs (Carroll et al., 1996; Eisner et al.,
1995; Mann and Plummer, 1995).
Biodiversity and Infectious Diseases
A central question in the field of biodiversity and conservation biology is whether increased diversity of host species tends to either buffer or
amplify disease outbreaks (Dobson, 2004; Dobson, 2005a). In the simplest
case the potential for disease outbreaks will be determined by the magnitude of R0 – this number is formerly defined as the number of secondary
infections produced by the initial infectious individual introduced into a
population of susceptible hosts (Anderson, 1982; Anderson and May,
1986; Dietz, 1993).
We can think of R0 as a heuristic mathematical device that allows us
to examine the biological conditions that lead to an outbreak; quantifying
the relative contribution of the ecological factors that determine the magnitude of R0 considerably focuses our understanding of how different
pathogens might be controlled. In cases where increases in species diversity lead to increases in the number of contacts between infected individuals and potentially susceptible hosts, increased host diversity will always
lead to increased values of R0 and a greater potential for disease outbreaks (Figure 3, see page 409). In contrast, where increases in inter-spe-
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cific transmission lead to reductions in within-species transmission, then
it is possible for increased host species diversity to lead to reductions in
R0. This will be the case for vector-transmitted pathogens such as malaria, yellow fever, dengue fever and Lyme disease (Dobson, 2004). On one
hand, increased host diversity will lead to increases in the resources available to the vector population (e.g., more blood meals). However, as most
vectors take a finite number of blood meals per lifetime, this will lead to
an increased proportion of bites wasted on hosts that may be less viable
resources for the pathogen (Figure 3, page 409). Whether or not the
pathogens are buffered or amplified by the increased host diversity will
depend on whether increases in the size of the vector population are sufficient to compensate for the ‘wasted bites’ on less viable hosts. Ecologists
have called this reduction in disease risk as host species diversity increases ‘the dilution effect’ (LoGuidice et al., 2003; Schmidt and Ostfeld, 2001).
In a world where global climate change may lead to range expansion of
vector transmitted pathogens from the tropics into the sub-tropics and
temperate zones, the dilution effect creates an important utilitarian argument for conserving biological diversity (particularly vertebrates) – as
long as these species are present in abundance, the biting rate of mosquitoes on humans should be reduced.
Rinderpest in the Serengeti
As a final example of the role that infectious diseases can play in modifying the structure of complex ecosystems, let us consider the impact
that a single introduced pathogen has had on species that live in the
African savanna; an ecosystem where a large proportion of the human
population are highly dependent upon a range of ecosystem services to
supply most of their food and economic well-being (Homewood and
Rodgers, 1991). The savannas of East Africa support a large pastoralist
population, their herds often share grazing habitat with wild antelope
species, and the millions of ecotourists that visit the region each year provide the major input into the local economy. Yet the region also provides
perhaps the best example of a pathogen completely modifying the structure of a food web: the introduction of the rinderpest virus into subSaharan Africa in the 1890s (Branagan and Hammond, 1965; Plowright,
1982). Rinderpest is a morbillivirus that infects hoofed animals: cattle,
wild buffalo, wildebeest, giraffe, and other large antelope. It is closely
related to both canine distemper (CDV) and measles, two of the com-
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monest diseases of humans and their domestic dogs; the recent evolution
of these three pathogens is intimately entwined with domestication of
dogs and cattle, this created the opportunities for the pathogen to establish in new host species, where a few mutations allowed it to differentiate
itself from rinderpest, which is the ancestral main trunk of the morbillivirus tree (Barrett, 1987). The split between the three pathogens is so
recent (<5000 years) that there is still strong cross immunity between
them; inoculation of dogs with rinderpest vaccine will protect them
against distemper. This again raises interesting questions about how we
classify pathogens in food webs where they may fail to establish a
dependence upon a host, but stimulate an immunological response that
allows the host to protect itself against invasion by a potentially lethal
natural enemy.
Rinderpest caused one of the largest pandemics in recorded history, it
took 10 years to spread from the Horn of Africa to the Cape of Good Hope,
during this time it reduced the abundance of many ungulate / artiodactyl
species by as much as 80% (Plowright, 1982). This in turn produced a transient glut of food for decomposers and scavengers, such as vultures and
jackals, however, this quickly lead to a massive reduction in food supply
for the predators that relied on wildebeest and other game for food. The
removal of the ungulates changed the grazing intensity on both shrubs and
grasses. This seems to have allowed some tree species to undergo a pulse
of recruitment, thus many of the fever trees that create woodlands in
damper areas of the savanna seem consist mainly of individual trees that
are now just over a hundred years old (Dobson, 1995; Dobson and Crawley,
1994; Prins and Weyerhaeuser, 1987). In contrast, reduced levels of grass
grazing led to an increased fire frequency, which prevented the establishment of miombo bushland that had previously covered the savanna. This
in turn modified the habitat for many of the predators that require thicker bush coverage to successfully attack their prey.
The development of a vaccine for rinderpest in the 1950s allowed
these processes to be reversed (Dobson, 1995; Plowright and Taylor,
1967). There is an instructive irony here: the presence of rinderpest in
wildlife was blamed as the major reason why it had proved almost impossible to establish large-scale cattle ranches in East Africa. The rinderpest
vaccine was largely developed to help the cattle industry, it was only ever
applied to cattle, but this in turn led to its disappearance from wildlife.
Thus cattle had been the reservoir and the repeated epidemics observed
in wildlife were in response to constant spillovers from cattle. Rinderpest
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vaccination has successfully eradicated the disease from most parts of
Africa, except in times of civil unrest, when declines in vaccination coverage allow it to resurge. The impact on wildlife has been spectacular: in
the Serengeti, wildebeest numbers have grown from around 250,000 to
over 1.5 million; buffalo have appeared in areas where they previously
unrecorded, and lion and hyena numbers have increased dramatically in
response to the enhanced food supply (Sinclair, 1979). This observation
strengthens our contention that predators are less effective than
pathogens in regulating host abundance. The numbers of some species
have declined, for example there are fewer Thompson’s gazelles perhaps
because of more competition for grassland forage, but more likely
because of increased predation pressure from the more numerous hyenas. African wild dogs have also declined since widescale rinderpest vaccination allowed their prey to increase in abundance; this may be primarily due to competition with hyenas, though there may also be increased
risk of infectious disease, particularly distemper (Creel and Creel, 1996;
Dobson and Hudson, 1986). In the absence of rinderpest, it may be that
wild dogs (and other carnivores such as lions), no longer acquire crossimmunity to distemper. The increased wildebeest abundance has reduced
the excess of dried grass during the dry season; this has in turn led to a
reduction in fire frequency, which has allowed the miombo bushland to
return to some areas of the Serengeti (Sinclair and Arcese, 1995; Sinclair
and Norton-Griffiths, 1979).
The key points to take from this example is that while the biomass of
the rinderpest virus in the Serengeti was always less than 10 kg, the virus
had an impact on the abundance and dynamics of nearly all of the dominant plant and animal species in the system – even those which it did not
infect. It strengthens the notion that indirect effects may be as important
in shaping species abundance and web linkages as are direct interactions.
It also reinforces the earlier comments that pathogens may be as powerful as predators in regulating abundance and distribution of free-living
species in natural ecosystems. Finally the rinderpest epidemic illustrates
the importance of considering not just infectious diseases of humans
when we examine the epidemiology of future human well-being. The pastoralist population of East Africa was heavily impacted by the great
rinderpest epidemic, as their herds died, the pastoralists either starved, or
changed lifestyles and changed from pastoralists to farmers who settled
and permanently farmed the same area for crops. In the absence of
wildlife, tsetse flies switched to feeding on humans and created a major
BIODIVERSITY AND INFECTIOUS DISEASE: WHY WE NEED NATURE
153
epidemic of sleeping sickness (Rogers and Randolph 1988); a vivid illustration of the dilution effect described above. The recovery of wildlife,
itself a direct consequence of the control of rinderpest in livestock, led to
the huge expansion of tourism which provides most of the financial input
into the local economy.
Outro
My friend’s Bible again crashes on the congressional committee room
table, ‘Extinction is a sin!’. We can now add that the changes in land use
that create endangered species also create bell-ringers for the decline of a
variety of economic goods and services that will become increasingly limiting. This creates a final irony. The most vociferous critics of environmental protection implicitly assume humans and their domestic livestock
will be amongst the last species to go extinct. This opinion is inherent in
the belief that human existence is independent of the welfare of other
species. It is a dangerous form of naivety. Our dependence upon other
species makes it highly unlikely that humans will be the last species left
alive on the planet. The simplest bet is that we will go extinct about
halfway down the list. With a bit of luck, and a deeper understanding of
food webs, our technological skills may delay this inevitable demise. This
is less naïve, but still assumes we can develop replacements for the services provided by biodiversity. The alternative is to conserve a significant
amount of biodiversity and explicitly acknowledge that Nature has a
value beyond our current ethical, economic and scientific understanding.
Ultimately, our greatest need from Nature may be the challenge it still
presents to human creativity. This is both beyond value and essential to
our long-term health and economic and spiritual welfare.
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